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Contaminants in the Subsurface: Source Zone Assessment and Remediation 4 Objectives for Source Remediation The remedial objectives of interest to the Army and other potentially responsible range from groundwater restoration and plume shrinkage and containment, to mass removal, risk reduction, and cost minimization. A realistic evaluation of the prospects for, or success of, a source remediation action requires the specification of these objectives with clarity and precision. The project manager and other stakeholders must know the full range of site remedial objectives, their relative priorities, and how they are defined operationally as specific metrics, in order to determine whether source remedial actions will contribute to meeting objectives for the site. The primary purpose of this chapter is to describe the many objectives possible at sites for which source remediation is a viable option, many of which have been institutionalized within regulatory, risk assessment, and economic frameworks for site cleanup. Failure to explicitly state remedial objectives appears to be a significant barrier to the use of source remediation. That is, the vagueness with which objectives for remedial projects are often specified can preclude effective decision making with regard to source remediation. Too often, either data presented on the effects of source remediation are irrelevant to the stated objectives of the remedial project, or the objectives are stated so imprecisely that it is impossible to assess whether source remediation contributes to achieving them. Evidence supporting the above was received by the committee over the course of its deliberations during numerous briefings on source remediation projects at Department of Defense (DoD) facilities and other sites, supported by extensive documentation on some of these remedial efforts. Other related documents were also reviewed, including many case studies on source remediation
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Contaminants in the Subsurface: Source Zone Assessment and Remediation efforts available at the U.S. Environmental Protection Agency’s (EPA) Technology Innovation Office. In many of the cases reviewed, remedial project managers (RPMs) appeared unable to articulate a clear a priori rationale for undertaking source remediation at a site or to quantify the extent to which source remediation efforts were contributing to accomplishment of remedial objectives. To a significant extent, these interrelated problems appeared to reflect the absence of unambiguously stated remedial objectives for the sites or of clear operational definitions of those objectives. For example, during a brief report on an attempt to use steam recovery to remove a source area from a relatively homogenous unconsolidated aquifer, the project manager expressed considerable frustration with the effort, not only from a technical point of view but, more important, from the point of view that it was consuming significant resources while not contributing to any reduction in human health risk. There was no complete exposure pathway to the contaminated groundwater at the site, nor any expectation of a complete pathway in the near future. This is illustrative of a situation in which an explicit operational statement of site objectives (e.g., a reduction of human health risk as estimated by the procedures specified in EPA’s Risk Assessment Guidance for Superfund, RAGS), if made prior to the attempt at source remediation at this site, might have led to a decision not to attempt source remediation. This widespread problem of vaguely formulated remedial objectives, tenuously linked to performance metrics, is neither specific to the issue of source remediation nor reflective of any unusual failure on the part of the specific project managers with whom the committee interacted. Rather, this ambiguity is embodied in long-standing national policy statements (i.e., the National Contingency Plan) and analytical procedures (as embodied in RAGS). It is compounded by the fact that multiple stakeholders at a site not only may have very different objectives, but may also use very similar language to describe those very different objectives. Moreover, a particular performance metric may potentially correspond to a variety of different objectives and accordingly be viewed quite differently by different stakeholders. Finally, both the DNAPL problem and the effects of source remediation efforts raise temporal issues that are very poorly addressed by conventional analytical frameworks for assessing risks to human health and the environment. This chapter describes a variety of substantive remedial objectives.1 It shows how a stated objective can be defined operationally by several different metrics 1 Substantive objectives are concerned with the results of a decision process, in terms of a physical change at the site. In contrast, procedural objectives focus not on the outcome of a remedial effort, but on the process by which a decision is reached (e.g., transparency of the decision process, opportunities for public participation). Procedural objectives are often as important to stakeholders as substantive objectives, but they are not considered here because they are outside the scope of this report.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation and how a particular metric may represent the operational definition of several different objectives. This complex relationship between ultimate cleanup objectives and the ways in which they are measured can be the source of considerable uncertainty in the evaluation of remedial alternatives. It can also mask serious differences in stakeholder priorities, which only become apparent when an apparently “successful” remediation fails to satisfy key stakeholders. This is not meant to imply that better specification of objectives and their relationship to metrics of remediation will ensure stakeholder satisfaction. An unambiguous delineation of objectives and metrics will, however, allow the decision on source remediation to be more clearly evaluated. The relevance of source remediation to different stakeholders’ objectives can be identified in advance, and progress can be measured. FORMULATING OBJECTIVES One source of ambiguity during site remediation is that various stakeholders may use similar (or identical) language to describe radically different objectives for remediation of a site. Thus, this section provides an approach to clearly describing stakeholder objectives. There are three critical, interdependent elements in the unambiguous specification of a remedial objective for a site: (1) identifying the objective, (2) determining the appropriate metric(s) to measure achievement of the objective, and (3) determining the status of the objective. Although difficulties in specifying each of these elements among the projects reviewed have been noted, the element of status is addressed first because it is often overlooked, followed by discussions of common objectives and the selection of appropriate metrics for an objective. Status of Remedial Objectives Status refers to the fact that any identified remedial objective can be seen either as important in itself or as a means to an end. In the former case, the objective is termed absolute or primary, while in the latter, the objective is functional. (For an exposition of the contrast between absolute and functional objectives, see Udo de Haes et al., 1996, and Barnthouse et al., 1997.) Consider, for example, the objective of reducing contaminant concentrations in groundwater to a specified level at a particular point in time and space. This may be mandated under a particular regulatory framework as a necessary feature of a successful remediation, in which case it represents an absolute objective. Failure to achieve these concentrations represents failure of the remedy. The identical criterion, however, could be selected as a means of ensuring that risks to human health have been reduced to an acceptable level. In this case, the objective is functional, because there may be other objectives that achieve a comparable degree of health protection, such as precluding use of contaminated groundwater. Confusion about whether an objective is absolute or functional is not uncommon
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Contaminants in the Subsurface: Source Zone Assessment and Remediation at a wide range of sites, particularly with regard to maximum contaminant levels (MCLs) and how they are viewed by various stakeholders. MCLs are frequently cited as an absolute regulatory objective. Indeed, MCLs (or non-zero maximum contaminant level goals, MCLGs) may be determined to be either an “Applicable” or a “Relevant and Appropriate” requirement under the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). However, they can also serve as a functional objective that supports an absolute statutory objective. For example, a state may have determined that all of its groundwater should be protected as a potential source of drinking water. Consistent with state law, a demonstration that concentrations were below MCLs would indicate that the groundwater resource had been adequately protected. The state might, however, be open to other indicators that its requirement of resource protection had been met.2 There is further complexity with the designation of status. For example, MCLs may serve as a functional objective supporting a higher-level functional objective of achieving an acceptable human health risk, which in turn serves the absolute objective of protecting human health. In this case, there are clearly alternative functional objectives, both to meet the risk assessment functional objective and to meet the absolute objective of protecting health. In the first case, it may be that the actual conditions of use would indicate an acceptable level of risk even if the MCL were exceeded. In the second case, there are any number of ways to interrupt the relevant exposure pathway (such as institutional controls or the provision of a public water supply). The distinction between absolute and functional objectives is important, because trade-offs among different absolute objectives cannot be accomplished at a technical level. Rather, they represent social value judgments that must be made among stakeholders.3 In contrast, trade-offs between functional objectives can be made at a technical level, subject to the requirement that equivalence in meeting the corresponding absolute objective can be demonstrated. Thus, functional objectives are fungible.4 In the above example, the project manager can achieve health protection by precluding use of the contaminated water or by lowering concentrations in groundwater. Similarly, but within the realm of physically specified objectives, if the absolute objective were defined as meeting a concentration at a specified point of compliance (e.g., a fenceline), the project manager could trade off between preventing contaminant migration from the 2 For example, higher concentrations of contaminants that would then be reduced during routine disinfection of raw water sources might be deemed acceptable. 3 A considerable body of literature on such judgments among qualitatively different environmental objectives has developed in the context of life cycle assessment (e.g., Udo de Haes et al., 1996). 4 “Fungible” refers to goods or commodities that are freely exchangeable for or replaceable by another of like nature or kind in the satisfaction of an obligation.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation source or capturing migrating contaminants before they reached the point of compliance. It is important to bear in mind that a given functional objective may serve more than one absolute objective, and also that a particular objective may be functional for one stakeholder and absolute for another. For example, limiting the migration of contaminants in groundwater beyond the boundaries of a site may serve the absolute objectives of meeting a state statutory requirement or preventing “chemical trespass” (Gregory, 1993). On the other hand, it may serve the higher-level functional objectives of limiting human health risk to an acceptable level (by reducing exposure potential) or avoiding the effort and uncertainty of applying for an alternative concentration limit under the Resource Conservation and Recovery Act (RCRA). Different stakeholders may all agree that limiting the migration of contaminants in groundwater beyond the boundaries of a site is an important objective, but they would likely have very different responses to any proposals for substituting an alternative objective. Metrics for Remedial Objectives Ultimately, accomplishment of (or progress toward) a remedial objective can only be evaluated if there is a measurable value or metric associated with that objective.5 Accordingly, any objective that cannot be stated directly in terms of a metric must be assigned one or more subsidiary functional objectives that can be formulated in terms of a metric. This is illustrated in the following simplified example. The absolute objective is the protection of human health. This is not directly measurable in most cases, where illness has not been recorded in a site-associated population. Accordingly, a common functional objective is the specification of a Hazard Index < 1.0 and a cancer risk estimate < 10–4 in a quantitative risk assessment. In practical terms, this requires a lower-level functional objective—that “exposure point” concentrations in groundwater be less than a certain value. Two alternatives can be employed to achieve this: (1) change contaminant concentrations at the defined exposure point or (2) change the exposure point, for example, by providing alternate water sources and prohibiting water use near the contaminant source. Each of these functional objectives has an associated metric—a revised concentration in the water at the relevant exposure point. It is important to bear in mind that not all metrics are as unambiguously specified as is the concentration of a particular chemical at a particular point in 5 Our use of “metric” differs from that of EPA (2003b), which describes classes of metrics that were not in fact measured but were inferred from measured quantities. In our use, such “Type II and Type III” metrics would be considered functional objectives.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation time or space, as will become clear in the following section, which discusses common objectives during source remediation and their associated metrics. COMMONLY USED OBJECTIVES Whether defined by the stakeholder as absolute or functional, there are a set of objectives that have been widely used in site remediation. In many cases examined by the committee, the identification of site objectives has been less than clear, such that metrics appropriate to one objective have been employed for a different objective to which they are not applicable. The following sections distinguish between alternative possible objectives (whether absolute or functional) in four areas. There are obviously other kinds of objectives dealt with at sites, including programmatic and societal concerns. Moreover, the list of objectives within each area is merely illustrative and far from exhaustive. The four areas are Objectives related to a physical change at the site Objectives related to risks to human health and the environment Objectives related to life-cycle and other costs Objectives related to the time required to reach particular milestones Physical Objectives There are a number of physical objectives that may drive the design and performance evaluation of source zone treatment methods. These include mass removal, concentration reduction, mass flux reduction, reduction of source migration potential, plume size reduction, and changes in toxicity or mobility of residuals. The specification of metrics for performance evaluation with respect to these objectives is typically easy since the objectives are related to physical, measurable properties. In some cases, however, considerable inference must be interposed between the available metrics and the physical objective. Each of these objectives and their associated metrics are described briefly below. Mass Removal Removal of contaminant mass from a source zone is a common objective at hazardous waste sites and may be either absolute or functional (depending on the stakeholders, the governing regulations, and other factors). Many of the source zone treatment technologies, particularly those that rely upon fluid flushing of the source area (including surfactant/cosolvent flushing, steam flushing, air sparging, and water flushing), are designed to remove contaminant mass from the swept zone. For these technologies, the injected fluids serve as a carrier medium to transport the contaminant mass to the surface. The mass removed is recovered at
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Contaminants in the Subsurface: Source Zone Assessment and Remediation the surface through the collection and treatment of these flushing fluids. Other source zone treatment methods, such as chemical oxidation/reduction, soil heating, or enhanced bioremediation, are designed to destroy or convert the form or phase of the contaminant mass in situ. Common metrics for the mass removal objective are the mass of contaminant recovered or destroyed and the percentage of the total contaminant mass present in the subsurface that was recovered or destroyed. The first metric is relatively straightforward for flushing technologies, for which mass removal is quantified by measuring the contaminant mass recovered as a component of the extracted flushing fluid. For destruction or conversion methods, however, mass removal is less easily quantified, and one must rely on indirect metrics. Under some conditions, the measurement of the concentrations of reaction by-products can facilitate inference of mass removed. Accurate mass balances, however, are typically difficult or impossible to achieve in such situations, since mass conversion or destruction methods do not rely upon injected fluid recovery. Finally, the ability to measure the percentage of the total contaminant mass in the subsurface that is recovered or destroyed depends on estimates of the total mass present, which, depending on site characterization data, may be quite poor. Concentration Reduction Even more common than the mass removal objective is the objective of reducing contaminant concentration within an affected medium (i.e., soil, sediment, groundwater, etc.) to a desired lower value. The obvious associated metric is contaminant concentration. Like mass removal, concentration reduction can serve as an absolute objective (e.g., meeting MCLs) or as a functional objective for reaching some other absolute objective (e.g., reducing exposure and consequently health risk). The use of reductions in contaminant concentrations as remediation objectives is common because regulations often specify concentration compliance levels. Concentration is defined as the mass of the target compound per volume (or mass) of the affected medium (pore water, core sample, solid sample). Thus, concentration compliance or target levels can be defined in a number of ways, depending on the sampled medium. For example, groundwater concentration is typically defined as the contaminant mass per volume of pore fluid, while solid-phase concentration is often defined as a contaminant mass per mass of the sampled solid phase. Although concentration is often viewed as a precise metric, it should be noted that it really represents an average value over the volume sampled. If the distribution of contaminant mass within a source zone is highly irregular, local concentrations within a source zone can vary substantially from one another and from the average concentration that would take the entire source volume into consideration. In this report, the term “local” implies that concentrations are sampled over small spatial intervals by extracting small volumes of pore
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Contaminants in the Subsurface: Source Zone Assessment and Remediation fluid, such that these concentrations are representative of a known physical location within the contaminated zone at the time of sampling. Remediation technologies seek to reduce concentration levels through mass removal, conversion, or destruction. Thus, it is important to note the connections between mass removal and concentration reduction as remediation objectives. Ideally, if the total mass of the contaminant were removed from the source zone, the concentration in the aqueous phase would be reduced to below detection limits. In practice, however, total mass removal is next to impossible, and may instead be confined to more “treatable” areas. Treatability will depend upon the selected remediation method, but it also depends on permeability (flushing potential), degree of aquifer material affinity for the contaminant, volume and distribution of NAPL present, and composition of the contaminant. For example, aquifer material with high organic content may tend to more strongly sorb contaminants, or the presence of a NAPL pool may limit the degree of contact between the contaminant and the injected flushing fluid. Due to subsurface heterogeneity, treatability typically varies spatially, resulting in a significant spatial variability in the distribution of contaminant mass within the source zone subsequent to treatment. Although some technologies (e.g., steam flushing) may be more robust in their mass removal behavior, it is expected that contaminant concentrations will be locally variable following treatment. Thus, it is very difficult to make generalizations about how removing a certain percentage of mass relates to achieving a certain percent reduction in contaminant concentrations. A number of field studies (e.g., Londergan et al., 2001; Abriola et al., 2003, 2005; EPA, 2003a) have documented that local concentration reductions are achievable with source zone remediation. Such concentration reductions have ranged over one half to two orders of magnitude. While it is possible that local concentration levels in the treatable zones may be substantially reduced by mass removal, local concentrations in less accessible DNAPL zones will likely remain high (at or near aqueous solubility levels) following treatment. Less accessible zones may also retain significant organic mass on the solid (through sorption) or within stagnant pore fluids (through diffusion). Furthermore, if local groundwater flow rates are low within the source zone under natural conditions, diffusion from less accessible zones may result in increasing contaminant concentration levels in the more accessible areas over time (often termed concentration rebound) once remediation operations have ceased. Thus, the potential effect of mass removal on local contaminant concentrations will be a complex function of source zone properties, including DNAPL distribution, natural groundwater gradients, and the spatial distribution of sorbed mass. Based upon these considerations, the use of local concentration within a source zone as a metric of remedial success is problematic, particularly in the absence of high-resolution sampling. More integrative metrics are discussed below.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation Mass Flux Reduction While measurements of local concentrations permit the development of a picture of the spatial distribution of contamination within the pore water in the source zone, mass flux quantifies the potential influence of these concentrations on a downstream receptor. Mass flux is typically defined at some cross-sectional (planar) surface selected downstream of the source zone and roughly perpendicular to the direction of flow. The mass flux at a particular location in this transect is defined as the mass of contaminant moving across the surface per unit area per unit time. The total mass flux (or, more accurately, mass flow rate) is then obtained by integrating the mass flux over the plane. The average mass flux can then be obtained by dividing the total mass flow rate by the area of the cross-sectional plane of interest (it should be noted that these approaches may not translate well to fractured flow systems). A related metric, the flux-averaged concentration, is determined by dividing the average mass flux by the average groundwater velocity in the cross section. Mass flux reduction can be a functional objective that may, for example, support the higher-order objectives of reducing exposure to downstream receptors or preventing the growth of a plume downstream of the source zone. Although conceptually attractive as a remediation objective, mass flux reduction is difficult to quantify in practice, as suggested in Chapter 3. Most existing methods typically involve measurement of contaminant concentration at distinct points in the selected transect. Transformation of these measurements to flux estimates requires application of assumptions about the groundwater velocities at the measurement points. Furthermore, computation of average fluxes from such measurements is subject to a high level of uncertainty. More integrative methods for estimating average mass flux are currently under development. These involve alteration of the flow field through downstream pumping or installation of in situ flow-through devices at selected downstream locations. The relationship between concentration reduction or mass removal and mass flux reduction is as yet poorly understood. Considerable research is currently being directed at developing information and methodologies for the prediction and quantification of mass flux reduction from data on source zone mass removal and on aquifer and contaminant characteristics. Box 4-1 illustrates some of these efforts. A more extensive numerical investigation of this sort suggests that a two-orders-of-magnitude mass flux reduction may be achievable following partial source zone mass removal in Type I media (Lemke et al., 2004). Reduction of Source Migration Potential Reducing the potential for the source to migrate into clean subsurface areas is a commonly stated objective of the projects reviewed by the committee. Many source zones are characterized by the presence of DNAPL pools, which are
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Contaminants in the Subsurface: Source Zone Assessment and Remediation BOX 4-1 Relationship between Partial Source Removal and Mass Flux Reduction Recent analytical and numerical modeling efforts provide evidence that partial source zone removal may result in significant (several orders of magnitude) reduction in posttreatment contaminant mass flux (Rao et al., 2002; Lemke and Abriola, 2003; Rao and Jawitz, 2003). Consider two simulated DNAPL cross-sectional saturation profiles shown in Figures 4-1(A) and 4-2(A). Here the modeled formation is based upon an unconfined sandy glacial outwash aquifer located in Oscoda, Michigan, at the site of a former dry cleaning business. Aquifer characterization efforts were conducted in support of a Surfactant Enhanced Aquifer Remediation (SEAR) pilot-scale test (see Chapter 5), designed to solubilize and recover residual tetrachloroethylene (PCE) from a suspected DNAPL source zone at the site (Bachman Road). As part of the SEAR design effort, alternative spatial variability models of the unconfined aquifer were developed from formation core data and were used to generate entrapped PCE distributions using the immiscible fluid flow model MVALOR (Abriola et al., 1992; Rathfelder and Abriola, 1998; Abriola et al., 2002). Further details pertaining to the spatial variability models and simulation conditions can be found in Lemke and Abriola (2003) and Lemke et al. (2004). The simulated spill involved the release of 96 liters of PCE over four grid cells at the top of the model domain at a constant flux of 0.24 liter·day–1 for 400 days, with an additional 330 days for subsequent organic infiltration and redistribution. Examination of the two saturation distributions in Figures 4-1 and 4-2 reveals that while both contain the same total volume of PCE, this mass is distributed more uniformly in Figure 4-1 than in Figure 4-2. Much of the mass in Figure 4-2 is contained within a thin pool, where saturations reach up to 91 percent of the pore space. Alternatively, in Figure 4-1 maximum saturations of PCE do not exceed 31 percent. The potential influence of source zone mass removal on DNAPL distributions is illustrated in Figures 4-1(B) and 4-2(B), which present saturation mass depletion profiles. Here PCE mass removal was simulated using a lab-validated version of MISER (Taylor et al., 2001; Rathfelder et al., 2001). The initial saturation profiles shown in Figures 4-1(A) and 4-2(A) were flushed with approximately 1.5 pore volumes of surfactant solution and 10 pore volumes of water. Further details pertaining to the simulated surfactant flush can be found in Lemke (2003) and Lemke et al. (2004). Inspection of Figures 4-1(B) and 4-2(B) reveals that different degrees of mass removal are predicted for the different initial mass distributions, even though each was subjected to the same flushing conditions. In Figure 4-1(B), flushing has resulted in 97.8 percent PCE removal, with the remaining PCE distributed in thin of pools of short lateral extent. The mass dissolution for the PCE distribution in Figure 4-2(A) is substantially less, with only 43.2 percent of the mass removed. Here much of the original pooled PCE persists, with maximum concentrations still ranging up to 86 percent (compared to 13 percent in Figure 4-1(B). Figures 4-1(c) and 4-2(c) illustrate the source zone PCE concentrations evolving from the saturation distributions shown in Figures 4-1(b) and 4-2(b). Notice that despite the substantial mass removal, concentrations are still quite high in the
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Contaminants in the Subsurface: Source Zone Assessment and Remediation FIGURE 4-1 PCE infiltration and subsequent mass removal behavior: (A) initial PCE saturation, (B) PCE saturation after surfactant flushing, and (C) aqueous PCE concentration after surfactant flushing. SOURCE: Adapted from Lemke and Abriola (2003) and Lemke et al. (2004).
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Contaminants in the Subsurface: Source Zone Assessment and Remediation (primarily local) human and ecological receptors. The risk assessment process integrates information on the physical conditions at the site, the nature and extent of contamination, the toxicological and physical–chemical characteristics of the contaminants, the current and future land use conditions, and the dose–response relationship between projected exposure levels and potential toxic effects (see Table 4-6 for toxicological data on DNAPL constituents and chemical explosives). The end result of human health and ecological risk assessment is a numerical value of potential additional risk to the hypothetical receptor from the contaminant source. The calculated risk values are compared to an acceptable target risk level or to a range of acceptable risk defined by the NCP or by state regulations. If the risk estimate is greater than the acceptable target risk level, target cleanup level objectives are identified for the site using the assumptions developed in the risk assessment related to potential levels of exposure. The overall purpose of risk assessment is to address the absolute objective of protecting human health and the environment. The risk assessment will determine if site-specific risk is above acceptable limits and the extent to which site risk needs to be reduced to meet the absolute objective. The risk assessment will also provide information that will support development of functional objectives, such as identifying which chemicals and exposure pathways contribute most to elevated risk. It will also help define metrics of success related to remediation. For example, the risk assessment may determine that levels of chlorinated solvents in groundwater would result in unacceptable risk if the groundwater is used as a source of potable water. If the ability to reduce the groundwater contaminant concentrations were limited by a lack of available technologies, the metric of risk reduction might be met by supplying an alternate source of potable water to residents. RAGS and the other risk-based methods commonly used to evaluate site risk and establish cleanup levels provide a standardized, systematic approach for estimating site risk. The standardized approach allows for relatively easy implementation of the methods at a large number of sites and allows sites to be prioritized for cleanup action. These methodologies and their strengths and weaknesses for different applications have been described and evaluated in other NRC reports (e.g., NRC, 1983, 1999b). Since all contaminated military facilities conduct their site investigations and cleanups under either RCRA or CERCLA (NRC, 2003), it is likely that at identified Army sites where source remediation is an option, a risk assessment has already been conducted or will be conducted in the future. Distinctions between Human Health and Ecological Risk Assessment The methods typically used in human health risk assessment are highly prescribed by RAGS and similar risk-based methodologies. RAGS requires that risk estimates for humans be protective of individuals and be based on the maximum exposure that is reasonably likely to occur. This risk estimate tends to be
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Contaminants in the Subsurface: Source Zone Assessment and Remediation conservative, that is, it is more likely to overestimate than underestimate true risk. Site-specific information concerning exposure can be used in calculating the risk estimate when the information is well documented. More typically, however, default exposure assumptions identified by the standardized risk-based methodologies and toxicity criteria developed by EPA based on laboratory animal testing data are used. Because of the uncertainly inherent in using animal data to predict toxicity in humans, the toxicity criteria recommended for use by EPA have incorporated modifying factors that result in far lower allowed chemical intake in humans. For practical reasons, the use of these standardized assumptions for exposure and toxicity in evaluating human health risk is encouraged by the regulators. The outcomes of this approach are relatively limited flexibility in accounting for site-specific conditions and risk estimates that represent higher-than-average exposure conditions. Box 4-4 further discusses the role of uncertainty in risk assessment calculations, and how uncertainty can be more quantitatively assessed in lieu of using the default assumptions discussed above. The ecological risk assessment process is far less prescribed in the published risk-based methodologies than the human health risk evaluation process for several reasons. In the ecological risk assessment process, ecological risk does not exist unless receptors and habitat are currently present at a site or are likely to be present in the future. Highly developed industrial sites are less likely to sustain ecological receptors and habitat. Unlike human health risk assessment, where risk to only one species is evaluated, ecological risk assessment must consider all ecological receptors present or potentially present in all environmental media potentially impacted. This evaluation requires a site-specific survey to determine what types of receptors (plants, animals, invertebrates, etc.) are present in each medium (soil, surface water, sediment, etc.). Finally, ecological risk assessment evaluates risk to the population of each species present, being concerned with risk to individual members of the population only when the receptor is classified as a threatened or endangered species by state or federal regulations. The available risk-based methodologies present a general framework for ecological risk evaluation (EPA, 1991b, 1992c, 1994, 1997b), but the type of evaluation conducted for a specific site is typically negotiated with the regulatory agency having responsibility for the site. Exposure Pathways at Army Facilities Explosives and DNAPL contamination at Army facilities can represent very long-term sources of contamination for soil, groundwater, and surface water. If the explosives or DNAPL contamination is present in relatively shallow soil (4–6 meters below ground surface), direct contact with contaminated soil (ingestion, dermal contact) could occur through or as a result of excavation activities that might bring contaminated subsurface soil to the surface. Army facility occupants and offsite occupants may indirectly contact contaminants in shallow soil
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Contaminants in the Subsurface: Source Zone Assessment and Remediation BOX 4-4 Evaluation of Variability and Uncertainty in Risk Assessment The inherent variability in exposure variables or population response and the lack of knowledge about specific parameters used in estimating risk can both affect the outcome of a risk assessment and the degree of confidence associated with the results. Evaluation of these sources of uncertainty is necessary to allow risk assessment results to be viewed in the appropriate context. “Variability” refers to the true heterogeneity or diversity that occurs within a population or a sample. Examples of factors that have associated variability include contaminant concentration in a medium (air, water, soil, etc.), differences in exposure frequencies or duration, or, in the case of ecological risk assessments, inter- and intraspecies variability in dose–response relationships. EPA risk assessment guidance (EPA, 1989) states that risk management decisions at Superfund sites will generally be based on an individual that has a reasonable maximum exposure (RME). The intent of the RME is to estimate a conservative exposure case (i.e., well above the average case) that is still within the range of possible exposures based on both quantitative information and professional judgment. In addition, EPA recommends conducting a central tendency exposure estimate (CTE), which is a measure of the mean or median exposure. The difference between the CTE and RME gives an initial impression of the degree of variability in exposure and risk between individuals in an exposed population (EPA, 2001a). If a risk assessment has been conducted using a point estimate approach, a range of point estimates can be developed to represent variability in exposures. To calculate RME risk estimates using this approach, EPA has developed recommended default exposure values to use as inputs to the risk equations (EPA, 1992a, 1996a, 1997b, 2001b). A CTE risk estimate is calculated using central estimates for each of the exposure variables, which are available from EPA guidance and other sources. For both RME and CTE risk estimates, site-specific data are used if they are available. The point estimate approach to risk assessment does not determine where the CTE or RME risk estimates lie within the risk distribution, and the likelihood that an estimated risk will be sustained cannot be determined. This leads to uncertainty as to what level of remedial action is justified or necessary. If a risk assessment has been conducted using probabilistic techniques, parameter distributions are used as inputs to the risk equations rather than single values. These distributions characterize the interindividual variability inherent in each of the exposure assumptions, and they are used with mathematical processes such as Monte Carlo simulation to estimate risk. The simulation output is a distribution of risks that would occur in the population, which provides a better understanding of where the CTE and RME risks occur in the distribution. A technique known as one-dimensional Monte Carlo analysis can be used to estimate the probability of occurrence associated with a particular risk level of concern (e.g., cancer risk of 10–6) (EPA, 2001a). Uncertainty is also inherent in every human health and ecological risk assessment because one’s knowledge of actual exposure conditions and receptor
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Contaminants in the Subsurface: Source Zone Assessment and Remediation response to chemical exposure is imprecise. The degree of uncertainty depends to a large extent on the amount and adequacy of the available facility-specific data. Typically, the most significant areas of uncertainty associated with receptor exposure include exposure pathway identification, exposure assumptions, assumptions of steady-state conditions, environmental chemical characterization, and modeling procedures. The toxicity values used in risk assessment must also be viewed in light of uncertainties and gaps in toxicological data. Information concerning the effect of a chemical on humans is often limited. Toxicity data are often based on data derived from high-dose studies using a specially bred homogeneous animal population. These data are extrapolated for use in predicting risk to a heterogeneous human population that is more likely to experience a low-level, long-term exposure (EPA, 2001a). Ideally, the uncertainty associated with each parameter used in the risk assessment would be carried through the evaluation process in order to characterize the uncertainty associated with the final risk estimates. However, since actual exposure conditions cannot be fully described, a variety of modeling strategies are available to evaluate uncertainty. If a risk assessment has been conducted using a point estimate approach, parameter uncertainty is usually addressed in a qualitative manner for most variables (EPA, 2001a). For example, the uncertainty section of a point estimate risk assessment document might note that soil sampling conducted may not be representative of overall contaminant concentrations and, as a result, the risk estimate may over- or underestimate actual risk. Uncertainty in the environmental concentration term is addressed quantitatively to a limited extent in a point estimate approach by using the 95 percent upper confidence limit (UCL) for the arithmetic mean concentration in the risk estimate, which accounts for uncertainty associated with environmental sampling and site characterization (EPA, 1992b, 1997c, 2001a). The 95 percent UCL is combined in the same risk calculation with various central tendency and high-end point estimates for other exposure factors. If a risk estimate is conducted using probabilistic methods, the uncertainty associated with the best estimate of the exposure or risk distribution can be quantitatively estimated using a two-dimensional Monte Carlo analysis. This analysis can provide a quantitative measure of the confidence in the fraction of the population with a risk exceeding a particular level. Additionally, the output from this analysis can provide a quantitative measure of the confidence in the risk estimate for a particular fraction of the population (EPA, 2001a). Compared to a point estimate risk assessment, a probabilistic risk assessment based on the same state of knowledge can provide a more complete characterization of variability in risk and a quantitative evaluation of uncertainty. In deciding whether a probabilistic assessment of risk should be performed, the key question is whether this type of analysis (vs. a point estimate assessment) is likely to provide information that will help in risk assessment decision making. To assist site managers in deciding what type of risk assessment is best suited to their site, decision-making tools such as a tiered approach developed by EPA based on “scientific management decision points” are available to help identify the complexity of analysis that may be needed (EPA, 2001a).
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Contaminants in the Subsurface: Source Zone Assessment and Remediation through inhalation of DNAPL contaminants that have volatilized and migrated to the ground surface. If groundwater contaminated with explosives or DNAPLs is used as a potable water source at an Army facility, exposure to facility occupants can occur through direct contact with the water during ingestion and via dermal contact. Indirect contact through inhalation of DNAPL volatile organic compounds that become airborne during water use or through migration to the ground surface or into occupied structures is also a possibility. The same type of exposure can occur offsite if contaminant migration has occurred or could occur in the future, and if groundwater is used by offsite residents as a source of potable water. Ecological receptors are most likely to contact contaminants from explosives or DNAPL after the contaminants have migrated through groundwater and have discharged to surface water. In these cases, the threat may be somewhat less given the dilution of the contaminant that is likely to occur once it is discharged to the surface water and given the rapid volatilization of many DNAPL contaminants to air. Ecological receptors are not likely to contact explosive or DNAPL contaminants in soil below the top meter unless excavation activities bring contaminated soil to the surface. The physical extent of the contamination and the timeframe required for its reduction to levels that represent an acceptable risk affect several elements of exposure assessment and subsequent risk characterization: The higher the concentration of the contaminant in the environmental medium, the higher the potential intensity of the exposure. The more widespread the source of contamination, the larger the potential population of receptors that may contact the contaminant and/or the higher the potential exposure frequency. The longer it takes to remove contamination from the environment, the longer the potential exposure duration. In many cases, these factors require that the overall objective of protecting human health and the environment be met through a combination of treatment and long-term site management actions. Time-Scale Considerations for Risk Assessment The risk-based methods typically used at contaminated sites evaluate carcinogenic and noncancer risk to a hypothetical individual over the course of the person’s lifetime. These methods do not factor the lifetime of a source of contamination into risk estimates. They do not typically evaluate the size of the population potentially at risk, nor do they consider risk beyond the lifetime of an individual (i.e., they do not consider cumulative risk to the entire population exposed for the lifetime of the source of contamination). These shortcomings are
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Contaminants in the Subsurface: Source Zone Assessment and Remediation serious, given that source zone cleanup may take decades to complete for technical and financial reasons, and some level of contamination is likely to be left in place9 for an extended period of time. Some types of chemical sources that represent particularly long-term problems in groundwater (e.g., chlorinated solvents) are known or are presumed to be highly toxic to humans. Only very low concentrations of these contaminants would be allowed in the groundwater if it were a source of drinking water. The timeframe required to achieve these low-level concentrations, either through natural site recovery or various remedial alternatives, may be so long as to be inconsistent with the timeframes implicit in risk-based methodologies. This can severely limit the ability of the risk assessor to differentiate what may be significant public health impacts, if they occur some time in the future. The RAGS model, for example, is a static examination of risk for a fixed population, assuming constant conditions for 30 years (40 years for a family farm) under Reasonable Maximum Exposure conditions. More conservative variants of this model may address full lifetime exposure. There are also more realistic models that address changes in contaminant concentration over time, as well as residential mobility, aging, and other demographic factors influencing exposure (e.g., Price et al., 1996; Wilson et al., 2001), but these also fail to address timeframes of contamination that may span centuries. Accordingly, existing risk metrics may be unable to demonstrate benefits from source remediation efforts, if the primary effect of those efforts is to reduce the time over which the source contributes to elevated contaminant concentrations in groundwater. If, hypothetically, a remedy were to have the effect on contaminant concentrations after an interval of several decades, it would not be detectable with current risk metrics. For the population that will reside in the area in the future, however, the risks from use of the groundwater have been substantially reduced. That is, in the absence of a remedy that will be effective within 30 years, existing analytical frameworks obscure important distinctions between remedies that are effective in 100 vs. 500 years. Techniques are available for longer-term types of risk evaluation; these techniques and associated models have been used for many years to evaluate risk associated with Department of Energy legacy waste sites where very long-lived radionuclides will be present in the environment for thousands of years (Yu et al., 1993; EPA, 1996b). Unfortunately, with chemical contaminants, estimation of population risk over the lifetime of the contaminant source is not typically conducted because there is no regulatory requirement to conduct such an evaluation, nor is there a currently prescribed regulatory context for considering the results of 9 “Contamination left in place,” as used in this report, is consistent with the interagency definition as hazardous substances, pollutants, or contaminants remaining at the site above levels that allow for unlimited use and unrestricted exposure (Air Force/Army/Navy/EPA, 1999).
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Contaminants in the Subsurface: Source Zone Assessment and Remediation such an evaluation. [It should be noted, however, that even though tools for evaluating long-lived contaminants are available, they are considered imperfect in predicting long-term risk because they rely on unverifiable assumptions about the future behavior of people and institutions (NRC, 2000a).] Existing risk assessment frameworks are badly in need of explicit reconsideration to better reflect the physical realities at sites if the best attainable remedies for these sites are to be selected. CONCLUSIONS AND RECOMMENDATIONS As mentioned in the opening of this chapter, clear definitions of absolute and functional objectives and metrics for success are not evident in most of the reports (both Army and non-Army) reviewed by the committee. This has made it difficult to determine the “success” of projects under any consistent definition. Reports of early (pre-2000) projects seldom contained sufficient rationale for how and why certain technologies were selected. More recent projects discuss objectives such as concentration reduction in the dissolved phase plume or reduction of source mass, but there is seldom evidence to suggest that the technology selected would meet those specific objectives. Indeed, within the Army several source remediation technologies have been piloted and then selected for scaling up in the absence of having specific cleanup objectives prior to the pilot projects. As an illustration of this, in situ chemical oxidation might have been attempted for a small portion of the source zone during a pilot study and found to achieve a certain percentage of mass removal. The committee observed that this would subsequently lead to full-scale implementation of the technology (1) without considering whether mass removal would meet the objectives of full-scale cleanup (which may be, for example, protection of human health) or (2) in the absence of any objectives for full-scale cleanup. Thus, in many cases observed by the committee, the decision to proceed with larger-scale remediation was not based on a demonstrated ability to achieve cleanup objectives. Rather, if the pilot test showed significant concentration reductions or mass removal, it was simply assumed that a larger-scale project would bring more widespread reductions. The following recommendations regarding objectives for source remediation are made. Remedial objectives should be laid out before deciding to attempt source remediation and selecting a particular technology. The committee observed that remedies are often implemented in the absence of clearly stated objectives, which are necessary to ensure that all stakeholders understand the basis of subsequent remediation decisions. Failure to state objectives in advance virtually guarantees stakeholder dissatisfaction and can lead to expensive and fruitless “mission creep” as alternative technologies are applied. This step is as important as accurately characterizing source zones at the site.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation A clear distinction between functional and absolute objectives is needed to evaluate options. If a given objective is merely a means by which an absolute objective is to be obtained (i.e., it is a functional objective), this should be made clear to all stakeholders. This is particularly important when there are alternative methods under consideration to achieve the absolute objectives, and when it is known or is likely that different stakeholders have a different willingness to substitute objectives for one another. Each objective should result in a metric; that is, a quantity that can be measured at the particular site in order to evaluate achievement of the objective. Objectives that lack metrics should be further specified in terms of subsidiary functional objectives that do have metrics. Furthermore, although decisions depend upon both technical and nontechnical factors, once a decision has been made, the focus should be on the technical metric to determine if remediation is successful. Objectives should strive to encompass the long time frames characteristic of many site cleanups that involve DNAPLs. In some existing frameworks, timeframes are very short (rarely longer than 30 years) relative to the persistence of DNAPL (up to centuries), such that alternative actions with significant differences in terms of the speed with which a site can be remedied cannot be distinguished. Within life cycle cost analysis, the chosen timeframe and discount rate can significantly affect cost estimations for different remedies. Decision tools with a more realistic temporal outlook have been developed in other areas of environmental science (e.g., storage and disposal of radioactive materials). Their application to DNAPL problems needs to be considered by the Army and by the site restoration community as a whole. REFERENCES Abriola, L. M., K. Rathfelder, M. Maiza, and S. Yadav. 1992. VALOR Code Version 1.0: A PC code for simulating immiscible contaminant transport in subsurface systems. TR-101018. Palo Alto, CA: Electric Power Research Institute. Abriola L. M., C. D. Drummond, E. J. Hahn, K. F. Hayes, T. C. G. Kibbey, L. D. Lemke, K. D. Pennell, E. A. Petrovskis, C. A. Ramsburg, and K. M. Rathfelder. 2005. Pilot-scale demonstration of surfactant-enhanced PCE solubilization at the Bachman Road site: (1) site characterization and test design. Environ. Sci. Technol. (In press). Abriola, L. M., C. D. Drummond, L. M. Lemke, K. M. Rathfelder, K. D. Pennell, E. Petrovskis, and G. Daniels. 2002. Surfactant enhanced aquifer remediation: application of mathematical models in the design and evaluation of a pilot-scale test. Pp. 303–310 In: Groundwater Quality: Natural and Enhanced Restoration of Groundwater Pollution. S. F. Thornton and S. E. Oswald (eds.). IAHS Publication No. 275. Wallingford, Oxfordshire, UK: International Association of Hydrological Sciences.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation Abriola, L. M., C. A. Ramsburg, K. D. Pennell, F. E. Löffler, M. Gamache, and E. A. Petrovskis. 2003. Post-treatment monitoring and biological activity at the Bachman road surfactant-enhanced aquifer remediation site. ACS preprints of extended abstracts 43:921–927. Adamson, D. T., J. M. McDade, and J. B. Hughes. 2003. Inoculation of a DNAPL source zone to initiate reductive dechlorination of PCE. Environ. Sci. Technol. 37:2525–2533. Air Force/Army/Navy/EPA. 1999. The Environmental Site Closeout Process Guide. Washington, DC: DoD and EPA. American Society for Testing and Materials. 1998. Standard Provisional Guide for Risk-Based Corrective Action (PS 104-98). Annual Book of ASTM Standards. West Conshohocken, PA: ASTM. Barnthouse, L., J. Fava, K. Humphreys, R. Hunt, L. Laibson, S. Noesen, J. Owens, J. Todd, B. Vigon, K. Weitz, and J. Young. 1997. Life-Cycle Impact Assessment: The State-of-the-Art. Pensacola, FL: SETAC Press. Department of the Army. 2002. Cost Analysis Manual. U.S. Army Cost and Economic Analysis Center. Environmental Protection Agency (EPA). 1986. Guidelines for Carcinogen Risk Assessment. Federal Register 51:33991. EPA. 1989. Risk Assessment Guidance for Superfund (RAGS): Volume I. Human Health Evaluation Manual (HHEM) (Part A, Baseline Risk Assessment). Interim Final. EPA/540/1-89/002. Washington, DC: Office of Emergency and Remedial Response. EPA. 1991a. Risk Assessment Guidance for Superfund (RAGS): Volume I-Human Health Evaluation Manual Supplemental Guidance: Standard Default Exposure Factors. Interim Final. OSWER Directive No. 9285.6-03. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1991b. Ecological Assessment of Superfund Sites: An Overview. Publication No. 9345.0-051. Washington, DC: US EPA Office of Solid Waste and Emergency Response. EPA. 1992a. Final Guidelines for Exposure Assessment. EPA/600/Z-92/001. Washington, DC: EPA. EPA. 1992b. Supplemental Guidance to RAGS: Calculating the Concentration Term. OSWER Directive No. 9285.7-081. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1992c. Developing a Work Scope for Ecological Assessments. Publication No. 9345.0-051. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1994. Field Studies for Ecological Risk Assessment. Publication No. 9345.0-051. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1995. Federal facility pollution prevention project analysis: a primer for applying life cycle and total cost assessment concepts. EPA 300-B-95-008. Washington, DC: EPA Office of Enforcement and Compliance Assurance. EPA. 1996a. Final Soil Screening Guidance: User’s Guide. EPA 540/R-96/018. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 1996b. Fact Sheet: Environmental Pathway Models-Ground-Water Modeling in Support of Remedial Decision Making at Sites Contaminated with Radioactive Material. EPA/540/F-94-024. Washington, DC: EPA Office of Radiation and Indoor Air. EPA. 1997a. Guiding Principles for Monte Carlo Analysis. EPA/630/R-97/001. Washington, DC: EPA. EPA. 1997b. Ecological Risk Assessment Guidance for Superfund: Process for Designing and Conducting Ecological Risk Assessments. Interim Final. EPA/540/R-97/006, OSWER Directive No. 9285.7-25. Edison, NJ: EPA Environmental Response Team. EPA. 1997c. Lognormal Distribution in Environmental Applications. EPA/600/R-97/006. Washington, DC: EPA Office of Research and Development and Office of Solid Waste and Emergency Response. EPA. 1998. Institutional Controls: A Reference Manual (Working Group Draft).
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Contaminants in the Subsurface: Source Zone Assessment and Remediation EPA. 1999. Department of Defense (DoD) Range Rule. Letter from Timothy Fields, Jr., Acting Assistant Administrator, U.S. Environmental Protection Agency, to Ms. Sherri Goodman, Deputy Under Secretary of Defense, Department of Defense, April 22, 1999. EPA. 2000. A Guide to Developing and Documenting Cost Estimates During the Feasibility Study. EPA 540-R-00-002. Washington, DC: U.S. Army Corps of Engineers and EPA Office of Emergency and Remedial Response. EPA. 2001a. Risk Assessment Guidance for Superfund: Volume III – Part A, Process for Conducting Probabilistic Risk Assessment. EPA 540-R-02-002, OSWER 9285.7-45. Washington, DC: EPA Office of Emergency and Remedial Response. EPA. 2001b. The Role of Screening-Level Risk Assessments and Refining Contaminants of Concern Baseline Risk Assessments. 12th Intermittent Bulletin, ECO Update Series. EPA 540/F-01/014. Washington, DC: EPA Office of Solid Waste and Emergency Response. EPA. 2002. 2002 Edition of the Drinking Water Standards and Health Advisories. EPA 822-R-02-038. Washington, DC: EPA Office of Water. EPA. 2003a. Abstracts of Remediation Case Studies, Volume 7. Washington, DC: EPA Federal Remediation Technologies Roundtable. EPA. 2003b. The DNAPL Remediation Challenge: Is There a Case for Source Depletion? EPA 600/ R-03/143. Washington, DC: EPA Office of Research and Development. Gregory, M. 1993. Health Effects of Hazardous Waste: An Environmentalist Perspective. Presented to ATSDR Hazardous Waste Conference in 1993. http://www.atsdr.cdc.gov/cx2b.html. Lemke, L. D. 2003. Influence of alternative spatial variability models on solute transport, DNAPL entrapment, and DNAPL recovery in a homogeneous, nonuniform sand aquifer. Ph.D. dissertation, University of Michigan, Civil and Environmental Engineering. Lemke, L. D., and L. M. Abriola. 2003. Predicting DNAPL entrapment and recovery: the influence of hydraulic property correlation. Stochastic Environmental Research and Risk Assessment. 17:408–418, doi 10.1007/s00477-003-0162-4. Lemke, L. D., L. M. Abriola, and J. R. Lang. 2004. DNAPL source zone remediation: influence of hydraulic property correlation on predicted source zone architecture, DNAPL recovery, and contaminant mass flux. Water Resources Research 40, W01511, doi:10.1029/2003WR001980. Li, X. D., and F. W. Schwartz. 2004. DNAPL remediation with in situ chemical oxidation using potassium permanganate: part I—mineralogy of Mn oxide and its dissolution in organic acids. Journal of Contaminant Hydrology 68:39–53. Londergan, J. T., H. W. Meinardus, P. E. Mariner, R. E. Jackson, C. L. Brown, V. Dwarakanath, G. A. Pope, J. S. Ginn, and S. Taffinder. 2001. DNAPL removal from a heterogeneous alluvial aquifer by surfactant-enhanced aquifer remediation. Ground Water Monitoring and Remediation 21:57–67. National Research Council (NRC). 1983. Risk Assessment in the Federal Government: Managing the Process. Washington, DC: National Academy Press. NRC. 1996. Understanding Risk. Washington, DC: National Academy Press. NRC. 1997. Innovations in Ground Water and Soil Cleanup. Washington, DC: National Academy Press. NRC. 1999a. Groundwater and Soil Cleanup: Improving Management of Persistent Contaminants. Washington, DC: National Academy Press. NRC. 1999b. Environmental Cleanup at Navy Facilities: Risk-Based Methods. Washington, DC: National Academy Press. NRC. 2000a. Long-Term Institutional Management of U.S. Department of Energy Legacy Waste Sites. Washington, DC: National Academy Press. NRC. 2000b. Natural Attenuation for Groundwater Remediation. Washington, DC: National Academy Press. NRC. 2003. Environmental Cleanup at Navy Facilities: Adaptive Site Management. Washington, DC: National Academies Press.
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Contaminants in the Subsurface: Source Zone Assessment and Remediation Nielsen, R. B., and J. D. Keasling. 1999. Reductive dechlorination of chlorinated ethene DNAPLS by a culture enriched from contaminated groundwater. Biotechnology and Bioengineering 62:160–165. Phelan, T. J., S. A. Bradford, D. M. O’Carroll, L. D. Lemke, and L. M. Abriola. 2004. Influence of textural and wettability variations on predictions of DNAPL persistence and plume development in saturated porous media. Advances in Water Resources 27(4):411–427. Price, P. S., C. L. Curry, P. E. Goodrum, M. N. Gray, J. I. McCrodden, N. W. Harrington, H. Carlson-Lynch, and R. E. Keenan. 1996. Monte Carlo modeling of time-dependent exposures using a Microexposure event approach. Risk Anal. 16(3):339–348. Rao, P. S. C., and J. W. Jawitz. 2003. Comment on “Steady state mass transfer from single-component dense nonaqueous phase liquids in uniform flow fields” by T. C. Sale and D. B. McWhorter. Water Resources Research 39:1068, doi: 10.1029/2001WR000599. Rao, P. S., J. W. Jawitz, G. C. Enfield, R. W. Falta, M. D. Annable, and L. A. Wood. 2002. Technology integration for contaminated site remediation: clean-up goals and performance criteria. Pp. 571–578 In: Groundwater Quality: Natural and Enhanced Restoration of Groundwater Pollution. S. F. Thornton and S. E. Oswald (eds.). IAHS Publication No. 275. Wallingford, Oxfordshire, UK: International Association of Hydrological Sciences. Rathfelder, K., and L. M. Abriola. 1998. On the influence of capillarity in the modeling of organic redistribution in two-phase systems. Advances in Water Resources 21(2):159–170. Rathfelder, K. M., L. M. Abriola, T. P. Taylor, and K. D. Pennell. 2001. Surfactant enhanced recovery of tetrachloroethylene from a porous medium containing low permeability lenses. II. Numerical simulation. Journal of Contaminant Hydrology 48:351–374. Sung, Y., K. M. Ritalahti, R. A. Sanford, J. W. Urbance, S. J. Flynn, J. M. Tiedje, and F. E. Löffler. 2003. Characterization of two tetrachloroethene-reducing, acetate-oxidizing anaerobic bacteria and their description as Desulfuromonas michignensis sp. nov. Applied and Environmental Microbiology 69:2964–2974. Taylor, T. P., K. D. Pennell, L. M. Abriola, and J. H. Dane. 2001. Surfactant enhanced recovery of tetrachloroethylene from a porous medium containing low permeability lenses. I. Experimental studies. Journal of Contaminant Hydrology 48:325–350 Udo de Haes, H., R. Heijungs, P. Hofstetter, G. Finnveden, O. Jolliet, P. Nichols, M. Hauschild, J. Potting, P. White, and E. Lindeijer. 1996. Towards a Methodology for Life Cycle Impact Assessment. Brussels: SETAC-Europe. Wilson, N., P. Price, and D. J. Paustenbach. 2001. An event-by-event probabilistic methodology for assessing the health risks of persistent chemicals in fish: a case study at the Palos Verdes Shelf. J. Toxicol. Environ. Health 62:595–642. Yang, Y., and P. L. McCarty. 2000. Biologically enhanced dissolution of tetrachloroethene DNAPL. Environ. Sci. Technol. 34:2979–2984. Yu, C., A. J. Zielen, J.-J. Cheng, Y. C. Yuan, L. G. Jones, D. J. LePoire, Y. Y. Wang, C. O. Lourenro, E. K. Gnanapragasam, E. Faillace, A. Wallo III, W. A. Williams, and H. Peterson. 1993. Manual for Implementing Residual Radioactive Material Guidelines Using RESRAD Version 5.0. ANL/EAD/LD-2. Argonne, IL: Argonne National Laboratory.
Representative terms from entire chapter: